Analyzing biomagnification of metals in different marine food webs using nitrogen isotopes

Marine Pollution Bulletin 56 (2008) 2082–2105

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Baseline

Edited by Bruce J. Richardson

The objective of BASELINE is to publish short communications on different aspects of pollution of the marine environment. Only those papers which clearly identify the quality of the data will be considered for publication. Contributors to Baseline should refer to ‘Baseline—The New Format and Content’ (Mar. Pollut. Bull. 42, 703–704).

Analyzing biomagnification of metals in different marine food webs using nitrogen isotopes Ma Shan Cheung, Wen-Xiong Wang * Atmospheric, Marine, Coastal Environment Program, Department of Biology, The Hong Kong University of Science and Technology (HKUST), Clearwater Bay, Kowloon, Hong Kong

High metal concentrations have been reported in different marine organisms (i.e., oysters and whelks; Blackmore, 2001), which may threaten the health of organisms higher in the food chain (Falco et al., 2006). Many studies have been conducted to investigate the bioaccumulation of metals in different marine systems (Reinfelder et al., 1998; Wang, 2002; Quinn et al., 2003), and generally indicated that biomagnification of metals in the marine food chains was not common. Gray (2002) concluded that methyl-mercury was the only metal which could be biomagnified, since many of the other metals were highly regulated and excreted by marine organisms. On the other hand, Wang (2002) suggested that metals such as Cd and Zn can potentially be biomagnified in marine systems, even though very few field studies showed biomagnification of metals in natural environments. To investigate the biomagnification of metals in nature, the trophic position of organisms in the food web needs to be known. Previously, gut contents were used to determine the diet and trophic position of an organism, however, this approach is time-consuming and labor-intensive. In addition, because different prey have different evacuation times in the gut, gut content analysis may underestimate the contribution of fast digested prey to the diet of some organisms and may fail to establish the trophic level of organism (MacDonald et al., 1982). Without a precise quantification of the actual trophic status in the food web, it is difficult to predict the relationship between the trophic level and the metal tissue concentration in organisms in nature. Nitrogen stable isotope ratios (15N) have now been employed to determine the trophic level of organisms. The d15N values provide a more accurate description of the trophic level occupied by a specific organism (DeNiro and Epstein, 1981; Lajtha and Michener, * Corresponding author. Tel.: +852 23587346; fax: +852 23581559. E-mail address: [email protected] (Wen-Xiong Wang).

1994; Post, 2002). The increase of d15N with increasing trophic level is mainly because the lighter isotope (14N) is more readily excreted. It is estimated that d15N will increase by an average of 3.4‰ per trophic level (Cabana and Rasmussen, 1994), although a wide range of d15N values is also reported (1.3–5.3‰, Minagawa and Wada, 1984). This technique has been used as a tool to investigate the relationship between the trophic level and the contaminant concentration in fish tissues (Kidd et al., 2001; Croteau et al., 2005). It can reveal the trophic structures and thus facilitate the interpretation of contaminant transport in aquatic communities. In this study, we combined the d15N and metal concentration measurements to study the metal food web dynamics in Hong Kong coastal waters. We employed the 15N technique to first elucidate the food web dynamics in Hong Kong rocky shores and soft bottom environment. We then coupled the food web dynamics with measurements of metal concentrations in different organisms to examine the relationship between the trophic level and the metal tissue concentration. One of the specific aims of this study was to examine whether biomagnification of a variety of metals (Cd, Cu, Ag, and Zn) occurred in specific food webs. We examined different benthic environments (soft bottom and rocky shores) to contrast the difference in metal biomagnification patterns from different food webs. To our knowledge, very few studies have employed the stable isotope technique to reveal the relationship between food web dynamics and metal biomagnification in aquatic food webs, especially for metals other than Hg (Stewart et al., 2004; Croteau et al., 2005). We collected samples from two rocky shore environments and one soft bottom environment. Clearwater Bay and Butterfly Bay were chosen as our sampling sites for the rocky shore environments due to their differences in hydrography (Fig. 1). Clearwater Bay is located at the eastern side of Hong Kong and characterized by high salinity throughout the year. This site can be considered

Baseline / Marine Pollution Bulletin 56 (2008) 2082–2105

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Fig. 1. Sampling sites at Clearwater Bay and Butterfly Bay from Hong Kong. Also shown is the benthic trawling site.

pristine since it is mainly impacted by the oceanic water from the South China Sea. Butterfly Bay is located at the western side of Hong Kong. With significant impact by the Pearl River Delta, the water from the western side of Hong Kong is characterized by low salinity in summer (due to large amount of freshwater input from the Pearl River) and high salinity in the winter (Morton and Morton, 1983). In addition, large amounts of organic matter and contaminants from the Pearl River Delta could also affect the organisms in the western region. For the soft bottom environment, we chose a site located at the western side of Lantau Island, Hong Kong. In addition to the influence from the Pearl River Delta, this site was also affected by the wastes produced during the construction of the Hong Kong International Airport. Similarly, this site was characterized by low salinity during summer and high salinity in winter. Seston and biofilm samples were collected from Clearwater Bay and Butterfly Bay in April 2005. For the seston samples, seawater was collected from the sampling site and stored in an acid washed Telfon bottle before it was transferred to the laboratory for filtration. The volume of seawater collected was adjusted by the total suspended solid in the water. In addition, the rocky shore organisms were collected from Clearwater Bay and Butterfly Bay on the same day. Macroalgae, Ulva sp. and Sargassum hemiphyllum were collected by hand during the low tide from Clearwater Bay, while only Ulva sp. was collected from Butterfly Bay because no other macroalgae was found. After collection, all macroalgae were carefully cleaned with seawater to remove any epibionts, and the samples were then placed in polyethylene bags before being transferred to the laboratory. The marcoinvertebrates were also collected by hand during the low tide. For both sites, several different species of invertebrates were collected, including limpets, barnacles, snails, and bivalves. All collected samples were stored at 80 °C until further analysis. The soft bottom samples were collected in August 2004 (summer) and February 2005 (winter) by Canton shrimp trawler during daytime. Trawling lasted for a total of 10 min at a speed of 3 knots h1. After trawling, the nets were lifted and samples were

collected from all the six nets and pooled together. The samples were first classified into different groups according to their families on board, and immediately stored at 4 °C before being transported back to the laboratory, where they were immediately placed at 80 °C until further processing. The seston were collected by filtering seawater collected from the sampling site on the acid cleaned glass fibre filters (GF/F). All the organisms were rinsed with ultrapure water and then dissected. All samples (except seston and biofilms) were freeze-dried and ground into fine powder using a mortar and pestle which were previously acid washed and ashed at 550 °C for 2 h. The tissue powders were divided into two groups for measurements of metal concentration and stable nitrogen isotope. Silver (Ag), cadmium (Cd), copper (Cu), and zinc (Zn) concentrations of the tissues were analyzed by inductively coupled plasmaatomic emission spectrometry (Perkin–Elmer, Wellesley, MA, USA) or atomic absorption spectroscopy (Z-8100, Hitachi, Tokyo, Japan). Briefly, tissues were dried in acid washed glass tubes at 80 °C until constant weight, and subsequently digested with 70% nitric acid at room temperature for two days under trace metals free condition, after which the samples were heated to 110 °C until totally digested. The metal concentrations were then measured after appropriate dilution and expressed on a dry weight basis. Random checks with standard reference materials (oyster tissue 1566b; National Institute of Standard and Technology, Gaitherburg, MD, USA) showed that recovery of metals was >90%. Nitrogen stable isotopes were analyzed using methods described in Quinn et al. (2003) with slight modification. Seston samples were treated with concentrated HCl to remove inorganic carbon in the samples and then dried at 60 °C for 48 h to remove the remaining HCl. The other samples were ground to a fine powder and lipids in the sample were extracted using methods described by Post (2002) (except seston and macroalgae which were analyzed for d15N values directly). Briefly, the powder samples were treated with 2:1 chloroform:methanol, and then centrifuged at 8000g. The pellets were soaked in 0.1 M HCl for 24 h to remove the remaining chloroform and methanol mixture. The

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Table 1 Concentrations (lg g1 dry wt) of Ag, Cd, Cu and Zn and d15N values (‰) in organisms collected from two intertidal rocky shores of Hong Kong Species

Group

Feeding type

Ag

Cd

Cu

Zn

d15N

Clearwater Bay Sargassum hemiphyllum Ulva sp. Septifer virgatus Nipponacmaea concinna Monodonta labio Saccostrea cucullata Tetraclita japonica Morula musiva Thais clavigera

Macroalga Macroalga Mussel (bivalve) Limpet (gastropod) Top shell (gastropod) Oyster (bivalve) Barnacle Gastropod Gastropod

– – Suspension feeder Grazer Grazer Suspension feeder Suspension feeder Predator Predator

0.03 ± 0.01 0.03 ± 0.02 0.02 ± 0.00 0.49 ± 0.22 0.41 ± 0.06 0.42 ± 0.17 0.21 ± 0.12 0.34 ± 0.24 0.51 ± 0.12

0.42 ± 0.09 0.03 ± 0.02 0.74 ± 0.19 1.14 ± 0.52 0.26 ± 0.05 0.73 ± 0.10 1.41 ± 0.47 2.35 ± 0.57 2.80 ± 1.94

2.90 ± 0.32 12.6 ± 0.91 6.78 ± 2.54 4.39 ± 1.04 79.2 ± 15.6 293 ± 36.4 3.06 ± 0.99 161 ± 64.6 357 ± 163

15.7 ± 2.42 15.0 ± 4.94 54.7 ± 13.5 72.9 ± 6.68 82.0 ± 5.30 1117 ± 316 524 ± 146 354 ± 126 237 ± 96.4

5.96 ± 0.34 7.41 ± 0.05 8.03 ± 0.15 8.28 ± 0.10 8.43 ± 0.30 9.72 ± 0.27 11.6 ± 0.21 11.7 ± 0.15 12.2 ± 0.23

Butterfly Bay Ulva sp. Planaxis sulcatus Monodonta labio Nerita albicilla Saccostrea cucullata Tetraclita japonica Nipponacmaea concinna Thais clavigera

Macroalga Gastropod Top shell (gastropod) Gastropod Oyster (bivalve) Barnacle Limpet (gastropod) Gastropod

– Grazer Grazer Grazer Suspension feeder Suspension feeder Grazer Predator

0.07 ± 0.01 1.80 ± 0.61 0.86 ± 0.34 0.43 ± 0.36 0.87 ± 0.38 5.51 ± 3.68 0.75 ± 0.20 2.19 ± 0.68

0.07 ± 0.06 0.15 ± 0.06 0.31 ± 0.33 0.62 ± 0.21 1.04 ± 0.32 1.15 ± 0.50 1.67 ± 1.79 1.77 ± 0.30

20.6 ± 11.4 82.4 ± 29.8 56.0 ± 22.4 18.8 ± 3.69 257 ± 82.9 9.18 ± 4.18 4.44 ± 2.49 203 ± 37.3

82.9 ± 45.0 94.6 ± 24.7 69.8 ± 2.27 123 ± 19.6 2108 ± 781 1965 ± 394 82.6 ± 19.0 366 ± 126

10.7 ± 0.06 12.4 ± 0.25 12.9 ± 0.15 13.1 ± 0.16 13.7 ± 0.11 13.8 ± 0.46 13.8 ± 0.22 17.3 ± 0.27

Data are mean ± SD.

samples were centrifuged again and the pellets were collected and dried at 60 °C. All the samples were packed in the tin capsules and analyzed for 14N and 15N by using DeltaPlusXL Isotope Ratio Mass Spectrometer (IRMS). All samples were standardized against atmospheric nitrogen as follows:

d15 N ð‰Þ ¼ ½ðRsample =Rstandard Þ  1  100 where the differential notation (R) represents the relative difference between isotopic ratios of the sample and standard gases (i.e., 15 N/14N). The d15N values of the seston and biofilm samples in Clearwater Bay were 4.08 and 1.14, respectively, and were depleted compared to samples from Butterfly Bay (12.59 and 11.79, respectively). The d15N values of the organisms collected from the rocky shore environments varied among species, from the highest value (17.3) in Thais clavigera (Butterfly Bay) to the lowest (5.96) in S. hemiphyllum (Clearwater Bay) (Table 1). The predatory gastropods (T. clavigera, Morula musiva) always had higher d15N than the other groups, indicating that they had higher trophic levels than the other species collected from the same sites. The macroalgae (S. hemiphyllum, Ulva sp.) had the lowest d15N values. Overall, the organisms may be grouped into different trophic levels based on the d15N values. For example, in the Clearwater Bay, the mussels Septifer virgatus, gastropods Nipponacmaea concinna and Monodonta labio may be grouped into the same trophic level due to their comparable d15N values. The barnacle Tetraclita japonica, and two predatory gastropods T. clavigera and M. musiva may also be grouped together. For the Butterfly Bay, the gastropods M. labio, Nerita albicilla, and N. concinna, the oyster Saccostrea cucullata, and the barnacle T. japonica had comparable d15N values. Maximum difference of the d15N values was 8.1 and 6.6 for the Clearwater Bay and Butterfly Bay, respectively, indicating that the length of food web in the Clearwater Bay was longer than that in the Butterfly Bay. Nitrogen isotopes (d15N) are a good indicator of trophic position of an organism, and the d15N values typically increase by 3–4‰ relative to its diet (DeNiro and Epstein, 1981; Cabana and Rasmussen, 1994; Kidd et al., 2001). The quantified d15N values appeared to realistically reflect the trophic positions in the Clearwater Bay rocky shores. For example, the d15N of seston was 4.08, and increased to 8.03 for the mussels S. virgatus, which fed exclusively on the seston particles (primarily the phytoplankton). The top predator T. clavigera had a d15N value of 12.2, which was about

3–4‰ higher than those measured in the mussels and herbivorous gastropods such as M. labio, the main diets of T. clavigera (Blackmore and Morton, 2002). T. clavigera collected from the Butterfly Bay also had the highest d15N value, which was 3–5‰ higher than its potential preys. Thus, the d15N data suggested that T. clavigera was the top predator on both rocky shore food webs (Clearwater Bay and Butterfly Bay). In addition, the d15N value of M. musiva from the Clearwater Bay rocky shore was comparable to T. clavigera, indicating that they had similar trophic position. However, except T. clavigera and M. musiva, the barnacles T. japonica also had a similar d15N value, showing that their trophic positions were similar. This can be explained by the feeding physiology of T. japonica. Different from the bivalves presented on the shore, barnacles also feed on larger food particles including zooplankton, thus their food web positions may be higher than the other purely herbivorous feeders. The d15N measured in the seston from Butterfly Bay was as high as 12.59‰, which was only slightly lower than those measured for suspension feeders such as oysters and barnacles. Indeed, by comparing the biofilm and seston samples from the two sites, we found that the d15N values were much higher in the samples collected from Butterfly Bay than the ones from Clearwater Bay. A similar trend was also observed in the organisms collected from these two sites. This enrichment of 15N can be explained by the local pollution in Hong Kong and human activity in the Pearl River Delta region. Elliott and Brush (2006) showed that with increasing population density and nutrient input from wastewater discharged to the watersheds, the d15N increased from +2‰ to +7‰. These wastewaters originated from human activity are generally enriched with d15N signatures. In addition, Moore and Suthers (2005) demonstrated that the livestock agriculture and residential development increased d15N values in the particulate organic matters of the estuaries. In our study, Butterfly Bay was affected by the local sewage outfalls and also strongly affected by the freshwater runoff from the Pearl River, and was subjected to inputs of fertilizers and organic wastes in the watersheds. Due to the enrichment of d15N in the watersheds, the stable isotope values in the organisms from Butterfly Bay also increased. In contrast, Clearwater Bay was mainly influenced by oceanic water and less subjected to nitrogen enrichment, and correspondingly, the d15N value was lower. Given the significant enrichment (and anthropogenic influence) of d15N signatures in the Butterfly Bay, the use of such technique to identify the trophic positions of animals may be more challenging.

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Baseline / Marine Pollution Bulletin 56 (2008) 2082–2105 Table 2 Concentrations (lg g1 dry wt) of Ag, Cd, Cu and Zn and d15N values (‰) in the organisms collected from soft bottom environment of Hong Kong at different seasons Zn

d15N

Species

Group

Feeding type

Ag

Cd

Cu

Summer Tonna sp. Perna viridis Scapharca sp. Turritella terebra Turricula nelliae spurius Bufonaria rana Nassarius crematus Pteroeides sparmanni Murex trappa

Gastropod Mussel (bivalve) Bivalve Gastropod Gastropod Gastropod Gastropod Sea anemone Gastropod

Predator Suspension feeder Suspension feeder Suspension feeder Predator Scavenger Scavenger Predator Scavenger

6.45 ± 0.35 0.15 ± 0.02 1.87 ± 0.02 1.81 ± 0.33 4.60 ± 1.08 2.55 ± 0.10 12.3 ± 0.62 0.01 ± 0.00 4.21 ± 3.22

2.37 ± 0.03 1.79 ± 0.35 13.2 ± 0.52 0.72 ± 0.11 5.83 ± 1.37 1.72 ± 0.02 4.66 ± 0.11 2.05 ± 0.51 10.4 ± 0.44

66.8 ± 0.60 12.4 ± 3.55 20.5 ± 0.87 192 ± 42.5 309 ± 67.0 49.6 ± 2.24 638 ± 44.8 5.94 ± 2.00 333 ± 15.4

629 ± 15.0 101 ± 17.1 232 ± 10.8 297 ± 59.0 221 ± 37.6 447 ± 18.1 469 ± 37.8 408 ± 138 274 ± 10.0

10.8 ± 0.12 11.2 ± 0.21 11.2 ± 0.07 13.9 ± 0.09 14.1 ± 0.08 14.1 ± 0.20 14.6 ± 0.18 14.8 ± 0.03 15.3 ± 0.06

Winter Cheilea sp. Anomia chinensis Potiarca pilula Ethusa indica Goniohellenus vadorum Clibanarius sp. Harpiosquilla harpax Bufonaria rana Scapharca sp. Nassarius crematus Charybdis acuta Charybdis japonica Parapenaeopsis hungerfordi

Gastropod Bivalve Bivalve Crab Crab Hermit crab Mantis shrimp Gastropod Bivalve Gastropod Crab Crab Prawn

Suspension feeder Suspension feeder Suspension feeder Predator Predator Predator/scavenger Predator Scavenger Suspension feeder Scavenger Predator Predator Predator

12.5 ± 0.47 3.38 ± 0.75 1.14 ± 0.09 0.63 ± 0.10 1.26 ± 0.02 0.85 ± 0.21 1.90 ± 0.14 4.81 ± 1.26 0.63 ± 0.01 12.0 ± 2.68 1.10 ± 0.01 0.48 ± 0.25 1.06 ± 0.05

0.66 ± 0.08 3.92 ± 0.98 6.29 ± 0.11 0.23 ± 0.31 0.20 ± 0.02 0.45 ± 0.10 1.10 ± 0.01 6.60 ± 1.60 6.20 ± 0.04 4.36 ± 0.71 0.20 ± 0.01 0.22 ± 0.07 0.41 ± 0.01

145 ± 2.42 83.2 ± 15.4 12.2 ± 1.84 38.3 ± 0.54 106 ± 1.36 291 ± 67.0 121 ± 2.56 57.8 ± 13.2 15.1 ± 0.18 534 ± 82.7 63.7 ± 0.21 39.4 ± 17.9 120 ± 3.69

89.3 ± 2.64 47.4 ± 11.3 178 ± 42.6 41.6 ± 12.0 43.0 ± 2.11 385 ± 91.0 103 ± 2.48 1477 ± 396 137 ± 2.75 580 ± 87.0 264 ± 5.45 125 ± 55.5 96.9 ± 2.48

7.24 ± 0.15 8.60 ± 0.35 10.5 ± 0.10 11.7 ± 0.45 12.1 ± 0.54 12.2 ± 0.20 12.9 ± 0.12 13.6 ± 0.06 14.4 ± 0.19 14.5 ± 0.11 15.0 ± 0.10 15.1 ± 0.05 16.8 ± 0.05

Data are mean ± SD.

In contrast, the d15N values in the soft bottom species did not show a clear pattern among different groups of organisms (ranging from 7.24 to 16.8, Table 2). The species composition in the soft bottom differed greatly between the two seasons of sampling. Only three species were simultaneously found during summer and winter (Scapharca sp., Bufonaria rana, Nassarius crematus). The metal concentrations in the seston collected from Clearwater Bay were 0.02, 3.12 and 134.5 lg g1 for Cd, Cu and Zn, respectively. For Butterfly Bay, the metal concentration in the seston was 0.10, 2.93 and 89.0 lg g1 for Cd, Cu and Zn, respectively. The Ag concentrations in seston from both locations were below the detection limit (0.1 lg g1). Organisms collected from the intertidal rocky shores of Butterfly Bay generally had higher Ag concentrations than the ones collected from Clearwater Bay (Table 1). For example, the Ag concentration in T. clavigera from Butterfly Bay was 2.19 lg g1 dry wt, which was 4 higher than that from Clearwater Bay (0.51 lg g1). For other organisms collected from Butterfly Bay, their Ag concentrations were about 2–10 times higher than the corresponding ones from Clearwater Bay. For Cd, Cu and Zn, their concentrations in corresponding species from the two sites were comparable. Among all the samples collected from Clearwater Bay and Butterfly Bay, the macroalgae had the lowest metal concentrations. The metal tissue concentrations varied among different taxonomic groups. In Clearwater Bay, T. clavigera had the highest Ag, Cd and Cu concentrations (0.51, 2.80 and 356 lg g1, respectively), whereas the oyster S. cucullata had the highest Zn concentration. In Butterfly Bay, the barnacle T. japonica had the highest Ag concentration (5.51 lg g1), and T. clavigera similarly had the highest Cd concentration among all the species (1.77 lg g1), whereas the highest Cu and Zn concentrations were found in S. cucullata (256 and 2108 lg g1, respectively). Metal concentrations varied among different phyla. Because different taxa accumulate metals via different routes and have different metal handling strategies, the correlations between metal concentrations in the environments and the whole tissue residues can be highly variable. The gastropods always had higher metal concentrations than the other groups, no matter in the rocky shores or soft bottom environment, especially for the predatory gastropods, T. clavigera and M. musiva. Such high accumulated me-

tal concentrations in the gastropods may be due to their high metal dietary assimilation efficiency and low efflux rate (Wang and Ke, 2002; Blackmore and Wang, 2004; Cheung and Wang, 2005). Besides the predatory gastropods, the oyster S. cucullata and the barnacles T. japonica also accumulated elevated concentrations of metals (especially Zn) in their bodies, consistent with many previous studies (Phillips and Rainbow, 1994; Robinson et al., 2005). Barnacles can accumulate very high Zn concentration in their bodies (e.g., 16,000 lg g1 in Balanus amphitrite, Rainbow and Smith, 1992). Higher metal concentrations (especially for Ag) were found in the organisms collected from the western side (estuarine water) than from the eastern side. Such elevation of metal concentrations may be due to the influences from the Pearl River. Ng and Wang (2005) reported that the dissolved Cd concentration in the western side of Hong Kong waters was 0.050–0.073 lg l1, as compared to 0.016–0.035 lg l1 in the eastern side. In addition to the anthropogenic input from the Pearl River discharge, the increase of metal concentration in the organisms from the western side may also be caused by the change of metal bioaccumulation kinetics due to a lower salinity. Ng and Wang (2005) showed that the increase in Cd uptake rate from the dissolved phase contributed to the higher Cd concentration in the green mussel Perna viridis from the western side of Hong Kong. For samples collected from the soft bottom environment, no noticeable difference was found between the metal concentrations in the same species collected in summer and winter (Table 2). Similar to the samples collected from the rocky shores, their metal tissue concentrations varied among taxa. Among all groups of organisms collected, gastropods always had higher metal concentrations when compared with organisms from other taxa. For example, N. crematus had the highest Ag (12.3 lg g1) and Cu (638 lg g1) concentrations, and Tonna sp. had the highest Zn (629 lg g1) concentration. Similar to the summer samples, the gastropods collected in the winter always had higher metal concentrations among all the taxa collected. Highest Ag concentration was found in Cheilea sp. and N. crematus, and highest Cd and Zn concentrations were found in B. rana. For the remaining taxa, their metal tissue concentrations were comparable. The large variation

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Butterfly Bay

Clearwater Bay 6

.8

Ag

Ag .6

4

.4 2 .2 0

0.0

Metal concentration (μg g

-1

dry weight)

6

3

Cd

Cd

4

y=-3.42+0.31x r2=0.685 p<0.01

2 y=-1.69+0.31x r2=0.710 p<0.01

2

1

0

0

600

400

Cu

Cu 300

400

200 200

100

0

0 4000

1600

Zn

Zn 1200

3000

800

2000

400

1000 0

0 4

6

8

10

12

14

10

12

14

16

18

δ15N value Fig. 2. Correlations between metal concentrations in organisms collected from rocky shores of Clearwater Bay (left panels) and Butterfly Bay (right panels) and their d15N values. Symbols for Clearwater Bay: (d) natural seston, (s) Sargassum hemiphyllum, (.) Ulva sp., (4) Septifer virgatus, (j) Nipponacmaea concinna, (h) Monodonta labio, () Saccostrea cucullata, (e) Tetraclita japonica, (N) Morula musiva and (5) Thais clavigera. For Butterfly Bay: (d) natural seston, (s) Ulva sp., (.) Planaxis sulcatus, (4) Monodonta labio, (j) Nerita albicilla, (h) Saccostrea cucullata, () Tetraclita japonica, (e) Nipponacmaea concinna and (N) Thais clavigera.

among different taxa could be explained by the differences in the metal bioaccumulation dynamics, as well as the species-specific metal exposure route. A significant positive correlation existed between the d15N and the Cd concentration in samples collected from Clearwater Bay (r2 = 0.710, p < 0.01) and Butterfly Bay (r2 = 0.685, p < 0.01) (Fig. 2). It is interesting to note that the slopes of the linear regression were comparable between the two locations. Cd concentrations increased from 0.03 lg g1 to 2.8 lg g1 and from 0.07 lg g 1 to 1.77 lg g1 along the food webs from Clearwater Bay and Butterfly Bay, respectively. In addition, there was a significant positive correlation between d15N and Cu in the samples collected from Clearwater Bay (r2 = 0.61, p < 0.05) when the barnacle data was removed. Cu concentration increased by 123 from 2.90 lg g1 to 357 lg g1 along the food web. No significant correlation was found between the d15N values and the concentrations of Ag and Zn from the two rocky shores (Fig. 2), or between d15N values and metals from the soft bottom environment in both seasons. In this study, using the combined d15N and metal concentration analysis in different groups of organisms, we demonstrated a sig-

nificant correlation between the d15N and the Cd concentration in samples collected from Clearwater Bay and Butterfly Bay. Such increase of metal concentration with increasing trophic level strongly suggested the biomagnifications of Cd in these specific food webs. Biomagnification is the transfer of a chemical from food to an organism, resulting in a generally higher concentration with increasing trophic level d15N (Gray, 2002), and can be inferred when there was a significant positive relationship between the metal concentration and the d15N value. Gray (2002) mentioned that most metals were not biomagnified due to the regulation and excretion during trophic transfer process. In our study, biomagnification was clearly metal-specific. We found that Cd was biomagnified in both rocky shore environments and Cu may be potentially biomagnified in Clearwater Bay. For Ag and Zn, no biomagnification was observed at the two rocky shore areas. For the soft bottom environments, no biomagnification of metal was also observed. Recently, Croteau et al. (2005) also found that Cd can be biomagnified in the freshwater lake system. However, there are very few field studies on metal biomagnification in different marine food webs. The biomagnification of Cd along the food webs may primarily be due to the unique Cd biodynamics in the top predatory

Baseline / Marine Pollution Bulletin 56 (2008) 2082–2105

gastropod. With a high metal assimilation efficiency and low metal efflux rate (Wang and Ke, 2002), predatory gastropods could accumulate high concentrations of Cd (Cheung et al., 2006). Since the slopes of the linear regression between the d15N and the Cd concentration in samples collected both rocky shores were similar, the dependence of Cd biomagnification on trophic level appeared to be comparable. However, the length of trophic levels was longer in the Clearwater Bay than in the Butterfly Bay, thus there was major difference of Cd biomagnifications between the two systems. In Clearwater Bay, Cd was biomagnified 93 times from the seston to the predatory gastropods, while in the Butterfly Bay, Cd was only biomagnified about 25 times. Such difference in the degree of Cd biomagnification between these two sites was mainly caused by the length of food web. In addition, the Butterfly Bay was subjected to significant anthropogenic influence, and the d15N values were significantly higher in the bottom of food webs than those found in Clearwater Bay, and varied by <2 times among different species collected from the same location (including seston). Blackmore and Morton (2002) demonstrated that the bioaccumulation of metals in the predatory gastropods, T. clavigera, could be influenced by their prey. The prey preference of the top predators from these two sites may be different, and may cause different bioaccumulation patterns. For Ag and Zn, they were not biomagnified in both rocky shore food webs. Ag has been shown to be mainly detoxified by forming insoluble granules (e.g., mussels and barnacles) and due to its low bioavailability, it may not be transferred efficiently along the food web (Wang, 2002). Zn was an essential metal which may be regulated by the bivalves, thus it may not be transferred along the food web with increasing concentrations. However, recent evidences suggested that Zn can be potentially biomagnified in marine juvenile fish, largely because of its higher dietary assimilation as compared to the larger size of fish (Zhang and Wang, 2007). In contrast to the rocky shore food web, no biomagnification was observed in the soft bottom environment when all the organisms were correlated with their stable isotopes values. Such absence of biomagnification from this environment may be due to the incomplete inclusion of the entire food web. Canton shrimp trawlers were used to collect the samples, thus only the bottom large invertebrates from the shallow sediments were collected. Other typical organisms such as polychaetes and oligochaetes were not collected, which are important prey for many benthic predators (Gray, 1981). Consequently, the correlation between the trophic position and the metal concentration may not be observed. In addition, trawling may mix up the organisms from different food webs, i.e., pelagic food web and soft bottom food web. Croteau et al. (2005) could not predict the biomagnification of Cd along the discrete epiphyte-based food webs. However, Cd enrichment along the food web was found after they separated the food web according to the class of animals. It was possible to observe biomagnification of metals in the benthic environment if the organisms were separated according to the food web. In summary, higher metal concentrations were found in the organisms from the western waters of Hong Kong due to anthropogenic influence and the lower salinity environments. The d15N signatures were also higher because of the extra input of organic matter from the Pearl River Delta. Our data demonstrated that biomagnification of Cd occurred in the intertidal rocky shore food web but not in the soft bottom food web. Biomagnification of metals was both food web and metal specific. The ability of metals being transferred along a food web was dependent on both the metal biokinetics of the top predator in the food web and the chemistry of the metals. When analyzing the correlation between metal tissue burden and trophic position of organisms from a complicated food web, unraveling the complexity of the food web is necessary. The stable isotope technique can provide a powerful tool to inves-

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tigate the relationship between metal trophic transfer and food web structure. Acknowledgements We thank Dr. I-Hsun Ni and Ms. Yi Ki Tam for their help in collecting the trawling samples, and one anonymous reviewer for his very thoughtful comments. This study was supported by a Competitive Earmarked Research Grant from the Hong Kong Research Grants Council (HKUST6420/06M) to W.-X. Wang. References Blackmore, G., 2001. Interspecific variation in heavy metal body concentrations in Hong Kong marine invertebrates. Environmental Pollution 114, 303–311. Blackmore, G., Morton, B., 2002. The influence of diet on comparative trace metal cadmium, copper and zinc accumulation in Thais clavigera (Gastropoda: Muricidae) preying on intertidal barnacles or mussels. Marine Pollution Bulletin 44, 870–876. Blackmore, G., Wang, W.-X., 2004. The transfer of cadmium, mercury, methylmercury, and zinc in an intertidal rocky shore food chain. Journal of Experimental Marine Biology and Ecology 307, 91–110. Cabana, G., Rasmussen, J.B., 1994. Modeling food-chain structure and contaminant bioaccumulation using stable nitrogen isotopes. Nature 372, 255–257. Cheung, M.S., Wang, W.-X., 2005. Influence of subcellular metal compartmentalization in different prey on the transfer of metals to a predatory gastropod. Marine Ecology Progress Series 286, 155–166. Cheung, M.S., Fok, E.M.W., Ng, T.Y.T., Yen, Y.F., Wang, W.-X., 2006. Subcellular cadmium distribution, accumulation, and toxicity in a predatory gastropod, Thais clavigera, fed different prey. Environmental Toxicology and Chemistry 25, 174–181. Croteau, M.N., Luoma, S.N., Stewart, A.R., 2005. Trophic transfer of metals along freshwater food chains: evidence of cadmium biomagnification in nature. Limnology and Oceanography 50, 1511–1519. DeNiro, M.J., Epstein, S., 1981. Influence of diet on the distribution of nitrogen in animals. Geochimica et Cosmochimica Acta 45, 341–351. Elliott, E.M., Brush, G.S., 2006. Sediment organic nitrogen isotopes in freshwater wetlands record long-term changes in watershed nitrogen source and land use. Environmental Science and Technology 40, 2910–2916. Falco, G., Llobet, J.M., Bocio, A., Domingo, J.L., 2006. Daily intake of arsenic, cadmium, mercury, and lead by consumption of edible marine species. Journal of Agricultural and Food Chemistry 54, 6106–6112. Gray, J.S., 1981. The Ecology of Marine Sediments: An Introduction to the Structure and Function of Bottom Communities. Cambridge University Press, Cambridge, UK. Gray, J.S., 2002. Biomagnification in marine systems: the perspective of an ecologist. Marine Pollution Bulletin 45, 46–52. Kidd, K.A., Bootsma, H.A., Hesslein, R.H., Muir, D.C.G., Hecky, R.E., 2001. Biomagnification of DDT through the benthic and pelagic food chains of Lake Malawi, East Africa: importance of trophic level and carbon source. Environmental Science and Technology 35, 14–20. Lajtha, K., Michener, R.H., 1994. Stable Isotopes in Ecology and Environmental Science. Blackwell Scientific Publications, Oxford, UK. MacDonald, J.S., Waiwood, K.G., Green, R.H., 1982. Rates of digestion of different prey in Atlantic Cod (Gadus morhua), ocean pout (Macrozoarces americanus), winter flounder (Pseudopleuronectes americanus) and American plaice (Hippoglossoides platessoides). Canadian Journal of Fisheries and Aquatic Sciences 39, 651–659. Minagawa, M., Wada, E., 1984. Stepwise enrichment of 15N along food chains: further evidence and the relation between 15N and animal age. Geochemica et Cosmochimica Acta 48, 1135–1140. Moore, S.K., Suthers, I.M., 2005. Can the nitrogen and carbon stable isotopes of the pygmy mussel, Xenostrobus securis, indicate catchment disturbance for estuaries in northern New South Wales, Australia? Estuaries 28, 714–725. Morton, B., Morton, J.E., 1983. The Sea Shore Ecology of Hong Kong. Hong Kong University Press, Hong Kong. Ng, T.Y.T., Wang, W.-X., 2005. Modeling of cadmium bioaccumulation in two populations of the green mussel Perna viridis. Environmental Toxicology and Chemistry 24, 2299–2305. Phillips, D.J.H., Rainbow, P.S., 1994. Biomonitoring of Trace Aquatic Contaminants. Chapman and Hall, London, UK. Post, D.M., 2002. Using stable isotopes to estimate trophic position: models, methods, and assumptions. Ecology 83, 703–718. Quinn, M.R., Feng, X.H., Folt, C.L., Chamberlain, C.P., 2003. Analyzing trophic transfer of metals in stream food chains using nitrogen isotopes. Science of the Total Environment 317, 73–89. Rainbow, P.S., Smith, B.D., 1992. Biomonitoring of Hong Kong coastal trace metals by barnacles, 1986–1989. In: Morton, B.S. (Ed.), The Marine Flora and Fauna of Hong Kong and Southern China III. Hong Kong University Press, Hong Kong, pp. 585–597. Reinfelder, J.R., Fisher, N.S., Luoma, S.N., Nichols, J.W., Wang, W.-X., 1998. Trace element trophic transfer in aquatic organisms: a critique of the kinetic model approach. Science of the Total Environment 219, 117–135.

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Robinson, W.A., Maher, W.A., Krikowa, F., Nell, J.A., Hand, R., 2005. The use of the oyster Saccostrea glomerata as a biomonitor of trace metal contamination: intrasample, local scale and temporal variability and its implications for biomonitoring. Journal of Environmental Monitoring 7, 208–223. Stewart, A.R., Luoma, S.N., Schlekat, C.E., Doblin, M.A., Hieb, K.A., 2004. Food chain pathway determines how selenium affects aquatic ecosystems. Environmental Science and Technology 38, 4519–4526.

Wang, W.-X., 2002. Interaction of trace metal and different marine food chain. Marine Ecology Progress Series 243, 295–309. Wang, W.-X., Ke, C., 2002. Dominance of dietary intake of cadmium and zinc by two marine predatory gastropods. Aquatic Toxicology 56, 153–165. Zhang, L., Wang, W.-X., 2007. Size dependence of potential metal biomagnification in a marine fish. Environmental Toxicology and Chemistry 26, 787–794.

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Organochlorine contaminants in the muscle of striped bass illegally harvested from shad gill nets in the Hudson River Estuary Ashok D. Deshpande a,*, Scott A. Doyle b, Bruce W. Dockum a, Amy Tesolin-Gee a,1 a b

National Marine Fisheries Service, Sandy Hook Laboratory, Highlands, NJ 07732, USA National Marine Fisheries Service, Law Enforcement, Brielle, NJ 08730, USA

From 1947 to 1977, the General Electric Company discharged into the Hudson River an estimated 2.7  105 kg of polychlorinated biphenyls (PCBs) from the transformer and electrical capacitor manufacturing facilities at Hudson Falls and Fort Edward, New York (Thomann et al., 1991). US Environmental Protection Agency (USEPA) estimates this PCB discharge to be 6.05  105 kg (USEPA 2002). Based on the documented purchase of 6.05  107 kg of PCBs, scientists believe that both PCB discharge values may be substantial underestimates (Skinner, personal communication). Sedimentladen contaminants have been intermittently reintroduced via high-discharge events and sediment disturbances caused by different dredging projects in the upper estuary. This has resulted in a downstream PCB gradient within the biota that is inversely correlated to the river kilometers (Rkm) away from the source (Bush et al., 1989; Limburg, 1986). In 1974, the US Environmental Protection Agency (EPA) Region II first documented the severity of PCB problems in the Hudson River (Nadeau and Davis, 1976). In fall 1975, the New York State Department of Environmental Conservation (NYSDEC) reported elevated levels of PCBs in the Hudson River fish. In 1976, NYSDEC banned all fishing in the Upper Hudson River and prohibited commercial harvest of most fish from Troy to New York City. In 1979, NYSDEC described the Hudson River as the most highly PCB contaminated river in the USA (Horn et al., 1979; Sloan et al., 2002). In 1985, the fishing advisories were extended to striped bass (Morone saxatilis) caught in marine waters and the commercial harvest of this species from marine waters was completely prohibited due to the PCB related human health concerns (New York State Department of Health, 1998). Rensselaer Polytechnic Institute (RPI) scientists have concisely summarized the chronology of PCB contamination of the Hudson River (RPI, 2003). American shad (Alosa sapidissima) ingress the Hudson River Estuary only for a limited time each spring to spawn, and as they eat very little during this residence, they are not expected to bioaccumulate significant amounts of PCBs from the Hudson River habitats (Smith 1985; Sloan and Armstrong, 1988). In the NYSDEC’s 1999 collections, PCB levels were low in shad from both less contaminated as well as those from more contaminated stretches of * Corresponding author. Tel.: +1 732 872 3043; fax: +1 732 872 3088. E-mail address: [email protected] (A.D. Deshpande). 1 Present address: The Dow Chemical Company, Midland, Michigan 48674, USA.

the Hudson River (Sloan et al., 2002, 2005). PCB concentrations in these shad ranged from 0.16 to 1.35 ppm, and did not exceed the US Food and Drug Administration’s (FDA) tolerance level of 2.0 ppm for human consumption. PCB averages for groups of striped bass, as well as PCBs in individual striped bass, in the same study were less than 2.0 ppm in less contaminated parts of the Hudson River in the vicinity of George Washington Bridge. For more contaminated upriver locations, the PCB averages were less than 2.0 ppm in the muscle of striped bass. However, there were a number of individual striped bass samples from these locations with levels exceeding 2.0 ppm, primarily in the males. Given the confounding nature of migratory behavior of striped bass (Secor and Piccoli, 1996) and the randomly observed high levels of PCB residues, it is understandable that only shad are presently permitted to be commercially fished in the Hudson River between the George Washington Bridge and the Bear Mountain Bridge (Fig. 1). The Hudson River Estuary Management Advisory Committee (HREMAC) estimates an incidental bycatch of approximately 21,000 kg or approximately 6100 dead striped bass each season in the fixed-gill nets set to capture shad along the shores of the Hudson River (HREMAC, 1999; TAC, 2001). It is illegal to catch striped bass with a net and commercial sale or possession for sale of striped bass from the Hudson River is illegal. It is also illegal to ship striped bass from New Jersey to New York. Fishermen are required by law to remove the bycatch and release it back into the river. In a federal investigation, federal agents and New Jersey State officers witnessed and video taped illegal harvest of striped bass entangled in shad gill nets about one mile south of the George Washington Bridge on the New Jersey side with the intention of selling them. The fisherman who confessed and admitted his guilt stated that the time and effort required in the removal and discard of bycatch of shad fishery results in substantial economic impacts on the fishing communities. The bycatch of striped bass was clearly more marketable than the shad both in price, size, and quantity during the shad run in April and May. Thus, the fisherman supplemented his income by marking illegally harvested striped bass as shad and selling them to fish wholesalers, who subsequently sold them to secondary dealers, and ultimately to seafood restaurants to be served to the consumers. We believe that just the small amounts of female shad were sold due to the low economic value. The fisherman took measures to avoid apprehension, such as